production from organic waste sources in a Norwegian context
Simon Alexander Saxegård
Master in Industrial Ecology
Supervisor: Helge Brattebø, EPT
Department of Energy and Process Engineering Submission date: June 2015
Norwegian University of Science and Technology
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Preface
The report have been written as a study assignment in the Industrial Ecology master programme at the Norwegian University of Science and Technology (NTNU) and is considered a contributing part to the BIOTENMARE research project.
First and foremost I want to give a well-deserved special thanks to Carine Grossrieder for her invaluable tutoring in SimaPro 8 and her willingness to follow up and support, even though outside her job description. I also want to give special thanks to Raymond Jørgensen at Frevar and Tore Fløan at ECOPRO for good cooperation and their sincere wish to contribute to both qualitative information and general overview, particular for Norwegian conditions. I also want to thank them for receiving me and my colleges with open arms and for giving us guided tours of their respective biogas plants. I also want to thank Hellen Hamilton for nice conversations and her positive attitude towards aiding me when I needed help. A special thanks goes also to Fredrik E. Solberg for good discussions about model development, in particular for the MFA modelling, and as a friend I always can ask for help and support. At last I want to thank Helge Brattebø for good discussions with resulting fruitful supervision and that he has let me contribute in the BIOTENMARE project by allowing me to do research in a scientific field I love and have particular interest for.
Trondheim, 12 Jun 2015
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Simon Alexander Saxegård
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Abstract
The waste management sector have attained increasing focus towards minimising environmental impact, resource recovery and efficiency. EU now bans disposal of organic wastes, in order to reduce groundwater pollution and greenhouse gas emission. Alternative organic waste treatment methods have therefore been implemented and studied. Biogas production from anaerobic digestion is one such treatment method, which have shown promising results for this purpose.
Biogas production produces two high utility products, biogas and bioresidual, that can reduce the consumption of fossil fuel and mineral fertilizer, respectively. As such, the impact reduction from biogas production can have implication for several sectors. These include waste management, agriculture and fossil fuel consuming sectors, such as the transportation sector.
In this study the purpose is to assess the environmental impact of the value chain for organic waste treatment, implementing substrates from several sectors. The study includes collection of organic waste for application of recovered energy and nutrient recycling, focusing also on the system expansion possibilities. A MFA based LCA model was developed in SimaPro 8, to assess environmental effect of biogas production in Norway in relation to the most common treatment method of today. Energy and nutrient recovery rates are included in this assessment.
The results from this study confirms previous studies by reviling that biogas production and end product utilisation contribute to low or net negative global warming potential (GWP) and fossil depletion potential (FDP). These results are in relation to the product it substitutes.
In relation to the compared treatment alternative, incineration of organic waste and manure applied as fertilizer, further strengthens assumption that biogas production is a beneficial solution, with respect to environmental impacts. Results also point to increased energy recovery rates and possible increased nutrient recovery.
For other impact categories is the situation different. The result found in this study show increased impacts for both human toxicity potential (HTP) and terrestrial acidification potential (TAP) relative to substituted products. The same is true for biogas production compared to the common treatment method for HTP but with various results for TAP.
The conclusion is that biogas production is beneficial in terms of GWP and FDP. The main stressors causing GWP impacts for biogas production are fossil CO2 form transport of organic waste, biogenic CH4 from storage of bioresidual or biogas post treatment, as well as N2O form application of bioresidual. FDP occur due to crude oil extraction for fossil fuels in transportation. TAP is caused exclusively (97 -99%) by NH3 emission. HTP main source is the level of heavy metal (HM) in bioresidual when applied for agricultural purposes.
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Sammendrag
Avfall sektoren har fått større fokus de senere årene som en mulig kilde til å redusere miljøpåvirkninger, øke resurs gjenvinning samt effektivitet. Eu har nå forbudt deponering av organisk materiale for å redusere grunnvannsforurensning og klimagassutslipp. Alternative behandlings metoder har derfor blitt utredet og bygd for å dekke kravet om behandling av organisk avfall. Biogass produksjon fra anaerob forråtning er en av de behandlings metodene som har fått økt fokus de senere årene, og gjennom mange studier har vist seg å være en av de beste behandling metodene.
Biogass produksjon gir to sluttprodukter, biogass og biorest, som kan benyttes til å redusere forbruket av fossil energi samt kunstgjødsel. På grunn av dette kan biogass produksjon redusere miljøpåvirkninger fra flere sektorer som avfalls behandling, jordbruk samt fossilt intensive sektorer som transport.
Denne studien tar for seg ulike miljøpåvirkninger gjennom hele livssyklusen relatert til verdikjeden for organisk avfallsbehandling fra en rekke sektorer. Studien tar for seg organisk avfallsbehandling fra innsamling av organisk avfall substrat samt mulige bruks områder for biogass samt biorest. Modellen utviklet i SimaPro 8, for å gjennomføre denne LCA studien, er basert på prinsipper fra MFA metodikk.
Resultatene fra denne studien underbygger tidligere studier på området, som tilsier at biogass produksjon er en god strategi for å redusere miljøpåvirkninger fra blant annet avfallssektoren.
Biogassproduksjon fører til reduserte utslipp for både klimagassutslipp (GWP) samt forbruk av fossile kilder (FDP). Resultatene fra denne studien peker mot økt gjenvinningsgrad av energi og muligheter for økt næringsgjenvinning i forhold til referanse situasjonen.
Referansesituasjonen i Norge er forbrenning av organisk avfall, da med annen type avfall, hvor gjødselen fra jordbruket blir benyttet direkte til spredning uten annen behandling.
For andre miljøkategorier peker resultatene en annen retning. Menneskelig toksisitetspotensiale (HTP) har i denne studien vist til store økninger i alle biogass casene, både i forhold til substituerte produkter men i tillegg sammenlignet med alternativ behandling. Samme tendensen gjelder for forsurings potensial for landområder (TAP), men hvor forskjellen mellom referanse case og biogass produksjon er mindre og hvor biogass i enkelte sammen henger gir lavere påvirkning.
Hovedkonklusjonen er at biomasseproduksjon er fordelaktig for GWP og FDP. For GWP er fossil CO2 fra transport samt unngåtte utslipp den viktigste stressoren sammen med CH4 fra lagring av biorest samt etter behandling av biogass. N2O fra bruk av biorest som gjødsel er dent tredje største faktoren her. FDP er et resultat av uthenting av råolje til produksjon av drivstoff.
HTP kommer hovedsakelig fra tungmetaller i bioresten anvendt som gjødsel, mens TAP kommer i all hovedsak (97 - 99%) fra NH3 fra spredning av biorest.
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Nomenclature
Biogas A gases produced by bacteria due to decomposition of organic material Biomethane Biogas with a content of >97.5% CH4,
Bioresidual Remaining organic and inorganic solids after digestion Biofertilizer Bioresidual applied as fertiliser
CBG Compressed biogas
CHP Combine Heat and Power
CH4 Methane
CO2 Carbon dioxide
Bioresidual The undigested leftover after anaerobic digestion
GHG Greenhouse gas emission
HM Heavy metal
HTP Human toxicity potential GWP Global warming potential FDP Fossil depletion potential
kWh kilo Watt hour, unit of energy during one hour, 3.6 MJ / 1 kWh LBG Liquid biogas, >98.5% CH4
LCA Life cycle assessment
MJ Mega Joule, unit of energy effect
N2O Dinitrous monoxide
NH3 Ammonia
Nm3 Normal cubic, one cubic of gas at 0 degrees Celsius MFA Material flow analysis
OMW Organic municipal waste
OIW Organic industrial waste
Sm3 Standard cubic, one cubic of gas at 15 degrees Celsius
SwSl Sewage sludge
TAP Terrestrial acidification potential
VS Volatile solids
Keywords: Life cycle assessment, Material flow analysis, biogas, biomethane, organic waste treatment, bioresidual, biofertilizer, organic waste incineration
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ix List of content
Preface ... i
Abstract ... iii
Sammendrag ... v
Nomenclature ... vii
List of figures ... xiii
List of equations ... xiii
List of tables ... xiii
1. Goal and scope ... 1
1.1. Background ... 1
1.2. Objective ... 2
1.3. Boundaries ... 2
1.4. Research question ... 3
1.5. Functional unit ... 3
2. Literature and empirical study ... 5
2.1. LCA of organic waste treatment and challenges ... 5
2.2. Previous LCAs of organic waste ... 6
2.3. Feedstock waste and effects on anaerobic processing ... 7
2.4. End product utility and substitution effects ... 12
2.5. Assessment of emissions and literature recommendations ... 16
2.6. Main treatment technologies ... 17
2.7. Anaerobic digestion ... 18
2.8. Incineration ... 19
2.9. Biogas upgrade technologies ... 20
2.10. Gas storage and application ... 22
2.11. Bioresidual end treatment ... 23
2.12. Norwegian empirical studies ... 23
3. Methodology ... 25
3.1. LCA ... 25
3.2. Lifecycle inventory description ... 26
3.3. Lifecycle impact assessment ... 26
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3.4. LCA aspects ... 27
3.5. Robustness of LCA results ... 28
3.6. Material flow analysis (MFA) ... 29
4. LCI ... 31
4.1. Choice of methodology ... 31
4.2. Data gathering procedure and critical assumptions ... 32
4.3. Model development ... 33
4.4. Feedstock estimation ... 35
4.5. Transport assessment ... 36
4.6. End-product approach and general calculations ... 37
4.7. Emission approach and general calculations ... 40
4.8. Sensitivity analysis development ... 44
4.9. Case sensitivity development ... 45
5. Results ... 49
5.1. LCA results... 49
5.2. MFA results ... 54
5.3. Parameter sensitivity ... 55
6. Discussion ... 57
6.1. Critical variables and process relationship ... 57
6.2. Main findings and agreement with literature ... 61
6.3. Uncertainty ... 62
6.4. Strengths and weaknesses of the method ... 64
6.5. Strengths and weaknesses of the model ... 64
6.6. Implications of this work ... 65
6.7. Challenges ... 66
6.8. Further work ... 66
7. Conclusion ... 69
References ... 71
Appendix 1 – Master thesis contract ... 77
Appendix 2 – Value chain for anaerobic digestion in Norway ... 80
Appendix 3 – Total impact results for sensitivity ... 81
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Appendix 4 – Sensitivity of different biogas upgrade alternatives... 82
Appendix 5 – Pasteurisation heat energy sensitivity ... 83
Appendix 6 – Sensitivity of different post treatment alternatives for bioresidual ... 84
Appendix 7 – Sensitivity of transport based on a 200% increase in distance for a single substrates and for the whole system, separately. ... 85
Appendix 8 – Sensitivity of biogas utility options ... 86
Appendix 9 – Substrate sensitivity ... 87
Appendix 10 – Substrate sensitivity GWP results ... 88
Appendix 11 – Bus sensitivity due to a change in persons per km ... 89
Appendix 12 – Nitrogen loss during storage and application of bioresidual and manure ... 90
Appendix 13 – NH3, N2O and N2 gas formation table REMOVE!!! ... 90
Appendix 14 – Net total NH3 emissions during storage and after field application ... 91
Appendix 15 - Net total CH4 emissions during storage and after field application... 91
Appendix 16 - Net total N2O emissions during storage and after field application ... 92
Appendix 17 –Process data for electricity requirement ... 92
Appendix 18 – Pasteurisation calculator ... 93
Appendix 19 – The process energy requirements for each case. ... 94
Appendix 20 - Main mass flows (simplified) ... 95
Appendix 21 – Inventory parameter list for case 1 and model functions from LCA model developed in Simapro 8 ... 96
Appendix 22 – Inventory calculation list from LCA model developed in Simapro 8 ... 127
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List of figures
Figure 1: The biochemical stages of anaerobic digestion... 9
Figure 2: LCA procedure and stages: (Baumann & Tillman 2004a) pg. 20. ... 25
Figure 3 Simplified value chain of anaerobic treatment ... 34
Figure 4: Net impacts for each impact category for each organic waste treatment case. Climate change is measured in kg CO2-equivalents(eq), Terrestrial acidification is measured in kg SO2-eq, Human toxicity is measured in kg 1.4 DB eq and fossil depletion is measured in kg oil eq. ... 50
Figure 5: Normalised results for all cases in comparison to all impact categories and process of origi ... 53
Figure 6: Relative parameter sensitivity results for Case 1 ... 56
List of equations
Equation 1: Total mass of waste analysed in the system determined by the DM (FU) ... 36Equation 2: Adjusted methane yield with respect to degrability ... 37
Equation 3: Basic methane equation ... 37
Equation 4: Methane equation with respect to increased biogas potential due to co-digestion ... 38
Equation 5: Methane content in biogas when mixing organic substrates ... 38
Equation 6: Biogas calculated on basis of methane produced ... 38
Equation 7: Total carbon dioxide produced in the AD ... 38
Equation 8: The mass of the biogas ... 39
Equation 9: Mass balance equation for bioresidual ... 39
Equation 10: Emissions caused by storage of digested organic wastes... 41
Equation 11: Nitrogen loss due to gaseous losses ... 42
List of tables
Table 1: Methane production from fat, proteins and carbohydrates ... 8Table 2: Properties of different waste ... 8
Table 3: Heavy metal concentration limitation for different bioresidual classes ... 13
Table 4: N based emission partition ... 16
Table 5: Inventory parameters for several biogas upgrading technologies ... 22
Table 6: Organic waste analysed in the Norwegian base cases ... 34
Table 7: Overall view of transport inventory and how end-product choice affects the GWP performance based on changes in tkm ... 37
Table 8: Emissions per type of bioresidual type relative to digested manure. Data based on Amon et al. (2006) ... 41
Table 10: Organic waste substrate composition for the sensitivity analysis of organic waste mix ... 44
Table 11: Parameter changes according to each case ... 48
Table 12: Stressor contribution GWP, positive emissions ... 51
Table 13: Stressor contribution GWP, negative impacts ... 51
xiv Table 14: Stressor contributor to caused HTP ... 52 Table 15: Total Transport required for each case ... 54 Table 16: Energy efficiency for each case with respect to output energy ... 54
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1. Goal and scope
In this chapter is this the background, objective and scope of the project described along with the contexts and purpose. A further explanation of the goal and scope will be assessed in chapter 3. This chapter is to give an introduction to the life cycle assessment (LCA) that have been performed in this study, and is an important part of any LCA study.
1.1. Background
Norway has a goal of reducing the environmental impact and have in this context, signed several agreements, such as the 2020 greenhouse gas reduction agreement, for restriction and reductions of pollutants. Agriculture, waste management and usage of fossil fuels are major contributors to the release of GHG’s such as CH4, N2O and CO2. A growing focus toward material and energy recycling have therefore been the focus of the last years politics, reaching a major breakthrough with EU’s ban on disposal of organic wastes to landfill.
Anaerobic treatment of organic wastes with biogas and biofertilizer production, have been analysed in several long term studies and found to be a highly recommendable treatment solution. Wastes, such as Manure, Organic Municipal Waste (OIW), Organic fats, Organic Industrial Waste (OIW) and Sewage Sludge (SwSl) are categorised organic waste substrates that anaerobic bacteria can digest and decomposed into biogas. Undigested volatile solids and inorganic solids are considered the dry fraction of the bioresidual, which can be applied as biofertilizer. Biofertilizer can thereby recycle nutrients and limit the use of mineral fertilizers (P) or artificial (N). Biogas can be used directly to produce heat and electricity or it can be upgraded to biomethane and LBG. These methane-purified gases have a higher level of utility than biogas due to a higher energy density and lower contamination of corrosive gases.
Biomethane is commonly used as fuel or can be mixed with natural gas in a remote grid. A downstream utilization of the end products from anaerobic digestion have the benefits of avoided emissions, thus resulting in avoided environmental impacts. For some impact categories however, can the biogas production value chain result in an increased level of stress, which is of interest to assess.
2 1.2. Objective
The objective in this study where to perform an LCA of biogas production and compare the treatment alternative with likely treatment options. The focus have been on environmental impacts such as (global warming potential, human toxicity, fossil depletion and terrestrial acidification, etc.) and how given assumptions and critical variables affect the environmental performance for a given treatment option. This study will focus on anaerobic treatment where biogas and bioresidual is utilized as substituting products in terms of common energy carriers and fertilizers, in a Norwegian context.
1.3. Boundaries
The goal of the project is to generate a model that can assess environmental performances of organic wastes in Norway. The study does not include any economic aspects or estimates. It is developed and analysed mainly as an LCA and any MFA is only complementary to identify the inventory required to perform the LCA. No specific case have been analysed, but several biogas plants have been used as inspiration for case development. Data have been collected from both state of the art literature and acquired information form empirical studies, executed in relation to this study, have been applied as sources. The mixed data have been applied both in the literature study and as foundation for the model development in SimaPro 8. This includes specific site data achieved through direct contact and earlier studies.
The analysis of the impacts is only midpoint characterized and any further impact
craterisation is not to be considered, this includes endpoint characterization and weighting of impact categories. The system is defined and only the most relevant flows have been assessed in the study, as figured in Appendix 2. Several organic waste substrates have been excluded because it is highly unlikely that any of these are to be applied as feedstock for biogas production in Norway. The system is to consider system expansion which special focus towards assessing life cycle substitution effects by applying biogas products and bioresidual for a wide spectre of commonly used products of today. It is also of high interest to
identifying the most common treatment options and assess how these affect the overall life cycle impact. All best available technologies (BAT) for storage and application of bioresidual have been excluded in this report.
3 1.4. Research question
The development of the research question and functional unit (FU) have been evaluated thought the methodology study. In this partial chapter is these defines an put in context to the study at hand.
To fulfil the goal and objective of this master have a research question been formulated. The purpose of the research question is to pinpoint the objective and goal, contributing to a higher degree of specification of the problem at hand. This study has been written in a Norwegian context and is focused towards critical variables that affect the environmental performance of this system. There is many aspect that have been included in this project such as a literature study, methodology study and model development. The most important is the model
development whereas the literature and methodology studies are only complementary to the model development itself. This along with the focus on environmental impacts generated from the treatment alternatives have made it clear that a parameter and variable analysis is the core of the project. The results and discussion are to highlight life cycle impacts of the waste treatment of organic substrates. This is to be performed in the context earlier described and to assess the variables responsible for causing or reducing the overall environmental impact including their process of origin. The conclusion are to summarise the most important findings from this study.
With the focus of interest, context and scope limitations have the research question been defined as:
“Which stressors are critically influencing the environmental life cycle impact of the biogas production in Norway, in comparison to the alternative organic waste treatment option, and which factors and variables limits or enhance these”
1.5. Functional unit
It was important that the functional unit enables the system to be comparable to major changes and still be comparable with the different results generated by the model developed in SimaPro 8, for each of the five cases. It was also, as described later in the methodology chapter, important that the research question is short and consistent. In this project there are several substrates that are treated and products being made. To be able to fulfil the goal and objective of this project a common constant had to be identified, something that remain the same in all cases for the model. Organic waste, have such a role in the model and have thus been chosen as the functional unit. It is described as follow: The treatment of one ton dry matter organic waste substrate.
The estimated energy and nutrient recovery as well as the environmental life cycle impacts will therefore be relative to this defined FU. Organic waste treatment contain large quantum of water, which in this FU is included by assessing the water content of each of the included
4 substrates, as explained later in this report. Nutrient and energy composition will be assessed in the same manner as the water content.
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2. Literature and empirical study
The literature study in this thesis considers previous LCA of organic wastes, technological options for organic waste treatment, in particular for anaerobic treatment and an empirical study of the treatment situation in Norway. The objective of this part of the study is to identify previous LCA studies and assess their findings to identify challenges concerning such studies and guidelines for data acquisition. There is a particular focus towards identifying the situation of anaerobic treatment and the alternative treatment method, in Norway particularly. Identifying important technologies and organic waste substrates, with respect to the parameter packages these represent is to be one of the main topics trough this review. In addition is possible utility purposes for biogas, bioresidual and for the produced heat from municipal incineration of organic wastes important to establish. In the end is a summary of the empirical data collected during field trips to Ecopro in Verdal and Frevar in Fredrikstad, presented.
2.1. LCA of organic waste treatment and challenges
LCA is a tool used to assess environmental impacts caused by the use or production of products, systems or services (Baumann & Tillman 2004a; Benoît & Mazijn 2009). The impact caused by these systems can be estimated with substantial certainty as they require a given set of input products, which can have predefined stressors caused by a given unit of use, the FU. Waste management scenarios pose a much higher degree of uncertainty. This is due to the fact waste is composed of a large variety of products that often are unknown and where the stressor inventory is much less defined. To solve this problem, an LCA of waste management requires data from various sources to satisfy a complete LCI. Data from various sources are often generic and does not give the exact waste composition for the case assessed. Therefore, the level of uncertainty is much higher in such studies and the results can often just be interpreted as close assumptions for the given case, thus not give a 100% correct impact estimate (Clavreul et al.
2012). It is therefore required that LCA studies are clear in which assumptions have been made and in which manner they have been acquired, particularly for the parameters that have a profound impact on the system.
The acquisition of LCI data can be defined in three major methodological approaches: (1) default variables, (2) theoretical technical data and (3) onsite-specific measurements (Clavreul et al. 2012). Default variables are highly generic data that cause a high level of uncertainty for a specific case, but give good generic data. Theoretical technical data describes processes that are defined by physical laws (natural gas law etc.). This gives the practitioner the possibility to calculate exact data, but it requires more work and a high level of detail. Onsite-specific measurements give highly reliable data that is relatively easy to access for a specific case. It is however, much less applicable in a generic context because every case is unique. Site measurements therefore lack the normal distributional that average generic data offers. It is also
6 important to be clear which method that have been applied in estimating the impact from the given LCI to diminish the uncertainty of methodological approach.
In LCA are there several methods to assess production systems and services of various complexity. This generates a source of uncertainty that have to be further assessed. For waste management this is of particular importance as waste is a commodity that has to be treated or disposed of in any case, and is thereby an inevitable process of any product. This makes it a necessity to compare waste management to a reference scenario so as to gain meaningful results (Clavreul et al. 2012). If not, the analysis is just a part in determining the impact caused by waste handling of a given product. For waste management LCA this is not the goal, but to assess the total impact of different treatment options of a homogenous waste mix. In conclusion it is important that the practitioner is aware of and transparent about the uncertainty level. Therefore an extra effort toward transparency for assumptions, sensitivity analysis and uncertainty propagation should be implemented in such studies (Clavreul et al. 2012).
2.2. Previous LCAs of organic waste
A wide variety of LCA’s have been applied to biogas production, anaerobic digestion and other organic waste treatment options the last decade.
The observed tendency, is that the actual impact caused by one or another treatment type highly varies dependent on the approach of study, area of interest and treatment option chosen for the given study. This makes it difficult to highlight specific results that are comparable between studies, but a general tendency towards a net benefit for several impact categories for anaerobic treatment compared to incineration has been witnessed (Poeschl et al. 2012a; Hamelin et al.
2014; Modahl et al. 2014; Khalid et al. 2011; Bernstad & Jansen 2011; Lyng et al. 2011).
Organic waste treatment is a subject of several treatment options, that in turn has a high degree of technological variation (Bernstad & Jansen 2011). This is especially true for anaerobic digestion treatment and biogas production systems where the end-product utility is an important part of the overall handling of the organic waste (Poeschl et al. 2012a; Modahl et al. 2014; Lyng et al. 2011). The end-products, biogas and bioresidual, has the possibility to substitute energy and fertilizer, respectively (Modahl et al. 2014). Such an assessment method of the end-products may generate negative net emissions due to the potentially saved impacts caused by the products substituted, such as mineral P and artificial N. For biogas can a multitude of energy carriers as fossil fuel commodities, natural gas in grid, remote heat or electricity in addition to other bio- fuels.
Several state of the art studies has concluded that anaerobic digestion options yield negative net GHG- emissions, but there is an increase in both nutrient enrichment and terrestrial acidification potential (TAP) compared to incineration (Bernstad & Jansen 2011; Lyng et al. 2011; Modahl et al. 2014). Other impact categories such as photochemical oxidant formation (POF),
7 particulate matter formation (PMF) and fossil depletion (FDP) contribute to net savings compared to the reference scenario, incineration (Hung & Solli 2012).
In the two continuous studies Poeschl et al. (2012a) and Poeschl et al. (2012b) several biogas plants were been assessed in a comprehensive LCA and the results were compared. Their results indicate that the biogas production is only yielding negative impacts for small biogas plants.
This study is performed in a German context where the German electricity mix is applied. For example heat requirements were covered by heat from natural gas. The study found positive impacts for GWP in the range of 191.68 kg CO2 eq for biomethane to natural gas substitution and 204.16 kg CO2 eq for fuel substitution. Only in the case where biomethane is applied for fuels cell application with 60% heat recovery are the GWP emissions net negative (-110.26). It was fond however that application of bioresidual results in net negative impact, for all cases.
The environmental impact is highly dependent on the type of waste being assessed in the feedstock, the study concludes. Their result is that the substrate yielding the most net negative impacts is straw. For cattle manure the result is – 23.22 kg CO2 eq, while municipal solid waste and slaughter house waste yields – 53.05 and 50.6 kg CO2 eq respectively.
2.3. Feedstock waste and effects on anaerobic processing
I this chapter, the relationship between organic waste substrates and anaerobic processing is presented. The findings comes from several state of the art studies, to secure complete data, eliminate epistemic1 uncertainty and system understanding in addition to the stochastic2 variations often found by surveying various studies.
Organic substrates and effects
The organic waste treated, highly determines the outcome of biogas production. This is mainly due to large variations in properties for each individual organic substrate (Carlsson & Uldal 2009; Poeschl et al. 2012a; Rehl & Müller 2011). Organic compounds consist mainly of the three organic carbons fat, protein and carbohydrates. These three, make out the volatile solid fraction (VS) of the organic substrate, table 1. The fraction that is not VS are considered ash weight (Carlsson & Uldal 2009).
1 Incomplete knowledge
2 Natural fluctuations or variations
8 Table 1: Methane production from fat, proteins and carbohydrates
Organic compound Biogas Methane Methane
Nm3/kg/ VS Nm3/kg VS %
Fat 1.37 0.96 70
Protein 0.64 0.51 80
Carbohydrates 0.84 0.42 50
Source: Extracted from Raadal & Morken (2008).
Many factors sustain, inhibits or increases the biogas production. The most notable of these factors are the alkalinity, VS, degrability, water content, nutrient composition, pH and temperature (Khalid et al. 2011; Søheim et al. 2010; Carlsson & Uldal 2009; Poeschl et al.
2012a; Hamelin et al. 2014; Bernstad & Jansen 2011; Raadal & Morken 2008; Jørgensen 2015;
Li et al. 2010). This indicates that the substrate mix is important when assessing biogas production and operation.
The processing option affects the biogas outcome of the anaerobic digestion to a certain extent, but overall is it the feedstock that determine the main range of the biogas and bioresidual composition and magnitude (Alvarez & Lidén 2008; Poeschl et al. 2012a; Carlsson & Uldal 2009).
In the report “Substrathåndboken” by Carlsson & Uldal 2009 experiments on biogas yield from a large variety of organic products (homogenously) tested where assessed and listed. Their results are show major variety in in digestion time, also called hydraulic retention time (HTR), methane yield and pH from the various products due to the substrate composition (Raadal &
Morken 2008; Schievano et al. 2011; Poeschl et al. 2012a), table 1. Many substrates contain organics, lignin or cellulose, that are difficult to digest, which thereby causes implications for the degrading of the VS, thus limiting the biogas production. This in turn, leads to a lower methane yield than normally assumed based on the VS content (Hamelin et al. 2014), table 2.
Table 2: Properties of different waste
Disaggregated substrates DMC (%)
Volatile Solids of DMC (%)
Degradability(D) of VS
CH4
%
Remaining solids (DS)9
CH4
yield m3/ ton VS1
Cattle manure 8%1 80%1 62%5 65%1 50.40% 213.01
Pig manure 8%1 80%1 62%5 65%1 50.40% 268.01
Frying fat 90%1 100%1 100%1 68%1 0% 757.01
Organic municipal waste
(OMW) 33%1 85%1 64%5 63%1 46% 461.01
Separated animal fats 4%1,2 95%1 100%7 60%7 5% 682.01
Fish processing waste
(offal) 42%1 98%1 65%6 71%1 36% 930.01
Sorted restaurant food waste 27%1 87%1 85%8 63%1 30% 461.01
9 Slaughterhouse waste
(blood) 10%1 95%1 65%6 63%1 38% 547.01
Slaughterhouse waste
(entrails) 30%3 83%3 63%3 63%3 48% 688.033
Slaughterhouse waste (offal) 16%1 83%1 65%6 68%1 46% 664.11
Diary processing 20%1 82%5 57%5 59%1 53% 277.05
Fruit and vegetable waste 15%1 95%1 57%6 60%7 46% 666.01
Sewage sludge 17%4 80%4 50%8 60%4 60% 336.04
Source: 1(Carlsson & Uldal 2009), 2(Gebauer & Eikebrokk 2006), 3 (Lyng et al. 2011) s. 22, 4(Wadahl 2014), 5 (Hamelin et al. 2014), 6Assumption based on (Hamelin et al. 2014), 7Other assumption, 8 (Sande et al. 2008). 9 Calculated as the remaining solids after digestion: 100% % ∗ % %
Anaerobic processing
Anaerobic digestion of organic material produces CH4 (50 – 80%), CO2 (15 – 40%), CO (0 – 0.3%), N2 (1 – 5%), NH3 (0 – 1%), O2 ( 0 – 0.5%) , H2 (0 – 0.3%), H2S (0.05 – 1.5%) by bacterial decomposition of organic material (Morken et al. 2007; Seadi 2002), fig 1. This process consists of the three main stages hydrolysis, fermentation, and methagonese, where fermentation can be described as two separate processes acidogenesis and acetogenesis, figure 1. The efficiency of the digestion besides degrability gradient is closely connected to four main parameters: Temperature, digestion time, pH and NH4+/NH3 concentration (Morken et al.
2007; Carlsson & Uldal 2009; Jørgensen 2015; Fløan 2015).
Figure 1: The biochemical stages of anaerobic digestion Figure derived from figure 2 in Morken et al. (2007)
10 Temperature
Temperature is one of the most important factors for biogas production, as it determines the rate at which the bioresidual decomposes and the biogas yield (Ariunbaatar et al. 2014; Yingjian et al. 2014; Raadal & Morken 2008; Carlsson & Uldal 2009). There are mainly three types of anaerobic digestion determined by the temperature occurring, both naturally and by human intervention. These are psychrophilic (7-25°C), mesophilic (25-42 °C) and thermophilic (49-60
°C) (Morken et al. 2007). As the temperature increase so does the methane yield, the degrability rate and the total process energy consumption (Morken et al. 2007; Mellbin 2010; Ariunbaatar et al. 2014).
In landfill is the presence of psychrophilic bacteria is the source for the CH4 and CO2 emissions emitted. Due to the long landfill time, the organic waste fully degrades and thus the resulting biogas are released directly into the atmosphere. Psychrophilic degradation has a hydraulic retention time (HRT) of 40d (Morken et al. 2007) and is thus little convenient at an industrial scale. Therefore, the major alternative options are either mesophilic (20-25d) or thermophilic (8-15d) (Angelidaki & Ellegaard 2003; Morken et al. 2007). The mesophilic alternative has historically been thought to be easier to start and maintain, in addition to require less energy than the thermophilic option. In later years has experiences shown that the thermophilic process are just as stable as the mesophilic, and thus deemed preferable du to the shorter HRT.
Ammonia and ammonium inhibition
The concentration of NH4+ in the digested material comes from several sources such as poultry manure, slaughter house waste offal and proteins, figure 1 (Carlsson & Uldal 2009; Morken et al. 2007). This affects the methane production mainly in two ways, by inhibiting hydrogen (H) to form bond with carbon (C) by forming NH3 (Morken et al. 2007) and by creating a toxic environment for the bacteria (Carlsson & Uldal 2009; Britto & Kronzucker 2002). Reversely works NH4+ and NH3 as a pH buffers as ammonia is converted, in the simplified reaction NH3+HNH4+, in an acidic environment, where responsively NH4+NH3+ H+ in a basic environment (Britto & Kronzucker 2002). This has also been the foundation of several NH3
inhibiting technologies used for bioresidual and manure storage (Bernstad & Jansen 2011;
Amon et al. 2006; Luostarinen et al. 2011). The toxic and inhibiting effect ammonium has on the anaerobic process disturbs the stability of the anaerobic process and is therefore an unwanted compound in the process (Ariunbaatar et al. 2014).
pH
The pH level are important for the process as it generates the conditions, in which the bacteria lives. Biogas production optimum occur at pH of 6.5 to 7.5 (Carlsson & Uldal 2009; Morken et al. 2007). During the hydrolysis decreases the pH, which is essential to maintain around the optimal pH 7. How low the pH reaches in the hydrolysis depends on the alkalinity of the
11 substrate in the process. The acidification occurs mainly due to increased CO2 levels in the substrate and the increasing presence of acids from proteins (Carlsson & Uldal 2009), figure 1.
To solve this problem can additives with a pH higher than 7.0, such as poultry manure or chemicals such as sodium bicarbonate (NaHCO3) and bicarbonate (HCO3) be used to increase the alkalinity. If the goal is to instantly increase the pH can Lye (sodium hydroxide, NaHO) be added to the process (Hauge 2014). Using Lye at a regular basis makes the process highly unbalanced and it will be difficult to maintain an optimal pH level over any length of time (Jørgensen 2015).
Co-digestion
There are many studies that describes the many benefits and often necessity for co-digestion of different substrates to maintain the biogas production (Carlsson & Uldal 2009; Raadal &
Morken 2008; Modahl et al. 2014; Lyng et al. 2011; Luostarinen et al. 2011). However, few has included a quantitative benefit assumption in their assessments because of the high uncertainty it poses.
The main benefits of co-digestion described in Carlsson & Uldal (2009) are pH buffering capacity, good mix of nutrients for the bacteria and a possibility for increased methane yield.
Co-digestion reduces or excludes the need for additives to the system such as micro nutrients, pH regulating chemicals (HCL and CaCH3) or buffering chemicals such as Ca(HCO3)2. In some instants, based on the water content of the feedstock, can virgin water consumption be reduced.
The more the bacteria manages themselves without the interference of human activity, the more stable and resilient the process becomes (Jørgensen 2015). Because of this, anaerobic digestion in Norway is in most cases, operated solely by the bacteria alone as described above. There is exceptions to this procedure and a homogenous feedstock has been, in all of these cases, the reason for those exceptions, (Jørgensen 2015). Manure slurry has proven to be a good co- digestion substrate due to its nutrient rich content water witch is essential for the anaerobic bacteria cultures (Labatut et al. 2011).
A quantitative co-digestion benefit of food waste and diary manure have been estimated to be in the range of 0.8 to 5.5 compared to digestion of manure alone (Li et al. 2010). A change in degrability has been the observed cause of this, which also affects the HRT during the acidification stage, figure 1. The optimal mixing ration observed is 6:1 food waste over manure and a acidification HRT of one day (Li et al. 2010).
Achieving a maximum biogas yield is a complex process that depends on the mix and type of input substrate and the nutrient ratio found within the bio-waste. Interestingly has a relationship between high amounts of water and an increase in methane yield been discovered, due to the solubility that CO2 has in water (Raadal & Morken 2008).
12 Biogas production, conclusion in literature
There have been observed a very complex biochemical system for biogas production and methane production yields, based on anaerobic bacterial produced biogas (Carlsson & Uldal 2009; Luostarinen et al. 2011; Bernstad & Jansen 2011; Morken et al. 2007). By having identified the most important parameters, it is possible to process organic waste anaerobically in an efficient way (Bernstad & Jansen 2011). There is however, not an easy task to predict the actual biogas production. In several studies, lower theoretical methane yield has been discovered than the actual been observed yield (Carlsson & Uldal 2009; Visvanathan 2014).
The main conclusion derived from these findings is that there is no absolute truth when it comes to methane yield theory, but that the theoretical potential is a guideline to assume the total production. To solve this problem, an onsite measurement is necessary to give any accurate information, where the result might prove to be somewhat misleading because of variations over time and from sample to sample. This requires special attention when trying to predict the output of methane based on the input substrate of choice (Carlsson & Uldal 2009) and should therefore be focused upon in the uncertainty assessment (Clavreul et al. 2012).
2.4. End product utility and substitution effects
The main driver for applying anaerobic digestion as a treatment alternative is the recycling potential of energy and nutrients the generated end-products offers (Raadal & Morken 2008;
Ariunbaatar et al. 2014; Lyng et al. 2011). There are mainly two by-products produced in the anaerobic digester, biogas and bioresidual. The bioresidual can be used for soil improvement or directly as fertilizer dependent on the heavy metal composition (Tormod Briseid 2010; Lyng et al. 2011; Luostarinen et al. 2011; Landbruks- og matdepatrementet, Klima- og miljødepartementet 2003). The bioresidual substitution effect are affected by several parameters such as nutrient content (N, P, Mg, K, C etc.), water content and the heavy metal content and the origin of the nutrient being substituted. Biogas, which is an energy carrier, can be applied for energy purposes an thusly substitute fossil or other common energy sources.
Bioresidual
Concentration of heavy metals are strictly regulated, both in EU and in Norway, and the greater the contamination the less utility and thus less substitution of conventional artificial or mineral fertilizers are possible. In Norway is it “Gjødselsforskriften” that covers the legalities for use of fertilizer and contamination limitation of bioresidual as a fertilizer (Landbruks- og matdepatrementet, Klima- og miljødepartementet 2003). The regulation distinguishes two forms of bioresidual, one from organic wastes and manure and one from or containing sewage sludge. From the point that the bioresidual is made or partly made of sewage sludge, the whole bioresidual body is treated as bioresidual from sewage sludge and a stricter use protocol is effective §15, 7. There are four levels of contamination in Norway, these are level 0, I, II and
13 III and are determined by the concentration of several heavy metals (HM) listed in table 3. If any of the HMs exceeds a class, the next class in effect even though the other HMs are within the restrictions levels several classes below.
The first level (0), allows the bioresidual to be used as fertilizer for root vegetables and fruits,
§ 24. Class 1 and 2 can be applied as fertilizer for surface dwelling crops. For bioresidual that contains sewage sludge there is special application procedure described in § 25. Sewage sludge containing biofertilizer cannot be applied for root vegetables and fruits because, these accumulates as substitutes for nutrients in plants and thus a higher ratio of unwanted bacteria and heavy metal contaminations might occur there. The bioresidual containing sewage sludge does also has to be ploughed into the soil within 18 hours after spreading, § 25, third sentence.
Table 3: Heavy metal concentration limitation for different bioresidual classes
Quality classes 0 I II III
mg/kg dry matter
Cadmium (Cd) 0.4 0.8 2 5
Lead (Pb) 40 60 80 200
Mercury (Hg) 0.2 0.6 3 5
Nickel (Ni) 20 30 50 80
Sink (Zn) 150 400 800 1500
Copper (Cu) 50 150 650 1000
Chromium (Cr) 50 60 100 150
Table replicated from Gjødselsforskriften §10 (Landbruks- og matdepatrementet, Klima- og miljødepartementet 2003).
When applying the bioresidual the presence of nutrients and the plant availability of these plays a crucial role in the substitution benefit (Luostarinen et al. 2011; Tonini et al. 2014). Luostarinen et al. 2011 has identified that the N in the manure alone are only 20 – 30% plant available.
Anaerobic digestion with some co-generation increases the N availability to 50 – 85%
(Luostarinen et al. 2011), but at the same time reduces the total amount of both nutrients, N and P (Amon et al. 2006; Möller & Müller 2012). For P is it assumed 100% availability, due to no data on the field (Möller & Müller 2012). The loses of N and P occur mainly due to losses by dewatering where the wet body is not applied as fertilizer, (Fløan 2015) in addition to the formation of NH3, N2O and N2 for the N losses (Bernstad & Jansen 2011). Loss of N and P is particularly severe if the bioresidual is separated and only one of the compartments, (wet or dry) is applied as fertilizer (Amon et al. 2006). The reason the loss is so extensive when only one of the bioresidual compartments are applied are due to differences in nutrient composition properties. 70% of N and 10% P found are in the wet body of the bioresidual (Poeschl et al.
2012a) and thus lost if not applied. Two studies, Möller & Müller (2012) and Rehl & Müller (2011), found that the average phosphors loss in the anaerobic digester is approximately 10%
14 of original P content. Why the phosphorus disappear is poorly documented and understood, but both their experiments resulted accordingly. A fair assumption for this losses could be due to retention in the digesters and storage tanks (Möller & Müller 2012). An aspect that is often forgotten when comparing the fertilization potential by application of biofertilizer is that it contains many more life essential minerals than just N and P. Commercial fertilizer often add N and P in addition to potassium (K) and the other minerals that is essential in smaller amounts is therefore not added such as sulphur (S) magnesium (Mg) etc. A net depletion of these minerals can lead to lower harvest yields, growth rate and quality of the food produced (Möller
& Müller 2012; Fløan 2015).
Alternative treatment of manure
Manure that would, in a Norwegian non-anaerobic digestion treatment context, otherwise been applied directly as fertilizer makes it important that the bioresidual is given back to the farmer.
Applying the bioresidual will continue to maintain the nutrient circle, in a greater extent than manure alone and secure that the farmers don’t suffer economically due to loss of fertilizer (Raadal & Morken 2008; Lyng et al. 2011; KLIF 2013).
Biogas
Biogas can be utilized for several purposes such as heat and electricity production, upgraded for gas grid supplement and fuel purposes. For biogas to be used in CHP no other treatment than cleaning is required (Lyng et al. 2011; Börjesson & Berglund 2006). Due to the relatively low concentration of methane in biogas (45 - 80%) is the gas often burned directly for energy purposes. If the production is greater than the demand for energy is the overproduced gas often torched (Fløan 2015). When the biogas is applied for grid purposes or as fuel is a upgrading in quality to biomethane (>97.5% CH4) required, which increases energy density and remove potential corrosive gases (Lyng et al. 2011; Luostarinen et al. 2011; Bauer et al. 2013; Hung &
Solli 2011; Hung & Solli 2012; Møyland 2012).
Emissions and impact categorization
For organic waste emissions are there several compartments that is of relevance. These compartments (air, soil, ground water, lake, rivers and marine environments) are effected in various ways by release of different stressors (Goedkoop et al. 2009).
Stressors has a tendency to contribute to only a few types of stress and it therefore measured in a given unit that best describes those impacts and is multiplied by a characterization factor for this given unit. This collection of stressors based on the effect is called an impact category, which is the categories in which LCA results are presented measured and interpret as the first step in an LCAI. For biogas production is the most relevant impact categories Global Warming Potential (GWP) given in kg CO2 eq, Terrestrial Acidification (AP) given in kg SO2 eq, human
15 toxicity potential (HTP) given in kg 1.4 DB-eq and fossil depletion (FDP) given in kg oil eq.
(Poeschl et al. 2012b; Modahl et al. 2014; Lyng et al. 2011). It has therefore been important to identify the stressor contribution to these impact categories.
Emission to air is caused by mainly the release of uncaptured CH4, CO2, NH3 and N2O from storage of manure and bioresidual (Luostarinen et al. 2011; Börjesson & Berglund 2006;
Andersen et al. 2010; Muha et al. 2014; Amon et al. 2006; Bernstad & Jansen 2011). During digestion itself there is measured low to no emissions of gasses (0 - 1% or 0.5 – 8% of produced biogas) because the treatment occurs in closed containers and pipes (Lyng et al. 2011; Jørgensen 2015; Fløan 2015). This is very case specific and is determined by the type of plant used. In Norwegian biogas plants is it most common to have water traps in the gas pipes so that the gas does not leak out of the system and as a result does these companies operate with no losses during digestion (Jørgensen 2015).
The two main emission stages, storage of manure and bioresidual is composed of as previously explained of two types of emissions, can be divided into C based and N based emissions (Muha et al. 2014). CH4 and CO2 is the two major C based gases that is released and both contributes mainly to the GWP category and is measured in terms of kg CO2 eq (Goedkoop et al. 2009).
NH3, N2O and N2 is the main N gasses produced but they contributes to different impact categories. NH3 has a great impact on TAP which is measured in kg SO2 eq (Goedkoop et al.
2009). N2O is a major contributor to GWP with a measured effect of 298 times the GWP of CO2 (Ecoinvent 2015). For biogenic CH4 is the characterisation factor 22.3 kg CO2 eq per kg.
In comparison, is the fossil CH4 25 kg CO2 eq per kg (Ecoinvent 2015).
N2 is a non reactive gas that makes up 79% of our atmosphere and has therefore no apparent impact, but contributes to losses of bioresidual mass and N fertilization potential (Möller &
Müller 2012; Bernstad & Jansen 2011). Emission such as those mentioned above occur in different stages of the organic waste treatment processing such as post storage or by spreading (Luostarinen et al. 2011; Amon et al. 2006; Modahl et al. 2014), Appendix 12 - 16. CH4
generation mainly occur due to remaining degradable material in the bioresidual, which is still under anaerobic condition (Amon et al. 2006; Bernstad & Jansen 2011). An aeriation of the bioresidual would therefore prevent unwanted post-produced biogas, Appendix 13. By covering the bioresidual with straw instead of wooden lid an increase of CH4 would occur, but at the same time reduce the generation of N2O and NH3 gases (Luostarinen et al. 2011). This effect occurs because more undigested carbon yields greater CH4 potential which the straw introduces (Amon et al. 2006). N2O and NH3 is mainly produced when the C/N ratio is high and added carbon reduces this value, and thus the potential NH3 and N2O formation (Bernstad & Jansen 2011). By digesting slurry the greatest benefit of methane capture and emission is evident, because of the carbon has already been transformed into biogas and thus the remaining C content is relatively low. As expected is the C/N ratio increased during anaerobic digestion and
16 the NH3 and N2O potential is thereby increased, resulting in greater emission rates of these gases if no inhibition measures are implemented (Bernstad & Jansen 2011; Luostarinen et al.
2011). As previously described, can a slight acidification of the bioresidual reduce the formation of NH3 and is among one of the most common inhibition methods applied. By using best available technology (BAT) can the overall reduction in N2O emission be up to 64%
(Luostarinen et al. 2011). For commercial N fertilizer can an emission rate of 2% of the total applied N might occur, but mainly determined by the pH of the soil it enters and poses therefore a high degree of uncertainty (Bernstad & Jansen 2011; Luostarinen et al. 2011). The production of N based gases varies a lot based on pH, C/N ratio, treatment option etc. The division of the N based gases has been found to vary drastically between different treatment technologies, table 4.
Table 4: N based emission partition
1Found in line 298 and 299 for NH3 and N2O respectively. 2 N2 is found by 100% - NH3% - N2O. 3 None of these sources where found, and therefore have the study that this is gathered from the main source. Source: Appendix 11 is just a short summary of table 5 in the paper (Bernstad & Jansen 2011), pg 1883.
2.5. Assessment of emissions and literature recommendations
The most common way to assess emissions in waste LCA is by applying transfer coefficient and emission rates found in the literature or by performing onsite measurements (Baumann &
Tillman 2004a; Carlsson & Uldal 2009). This is however, yielding unnecessary high uncertainties in a generic context (Clavreul et al. 2012). Muha et al. (2014), another study that assess the emissions from anaerobic treatment, has a rather blunt approach to the calculation of methane emissions caused in the storage tank. The CH4 emissions is calculated based on the methane yield (MY) per ton dry matter (DM) volatile solids (VS) with a subtraction of the amount already digested in the anaerobic process, determined by the degrability coefficient (Dg) and the average methane yield. Such an approach secures a mass balance correct estimation of emissions and the remaining bioresidual. The reliability of a study that has performed this approach should be higher, they argue, than a study based of predefined emission parameters. This is particularly important for co-digestion facilities because the amount of C and N vary much with the composition of the current feedstock batch in process. The same approach can to some extent be applied for N based emissions but a greater dependence on N loss from other sources is necessary and is thus a source of uncertainty.
Treatment type NH3 N2O N2 Sources in paper
AD 96%1 0.77%1 3.23%2 Chung (2007)3
Composting 2.40% 1.40% 96.20%
Sonesson (1996)3
17 Emission to soil is as previously explained caused by increasing levels in heavy metal concentration, but also the remaining levels of chemical oxygen demand (COD) and biological oxygen demand (BOD). High levels of these generates a anaerobic soil that is poor for life and has the potential to continue to produce biogas, which will be released directly to the atmosphere. A high biogas yield reduces the remaining COD and BOD and is therefore preferable in this context as well.
Emission to waterbodies3 comes in mainly two forms, leaching of nutrients from agricultural spreading and COD and BOD of the leached material. The nutrient N and P is worldwide a common problem due to agriculture and leaching from fields. It has been observed a big difference in leaching based on soil type the bioresidual is spread on (Bernstad & Jansen 2011).
Sandy soil has leaching coefficients of 25% to surface waters and 45% to ground water while loamy soil has 0% and 22% of sandy soil and loamy soil respectively for bioresidual.
2.6. Main treatment technologies
When considering waste treatment is there several options, both alternatives and technological options, for each of the technologies. Some of these technologies has been studied closer in this chapter to identify differences in emission, resource and energy requirements.
Pre-treatment
Pasteurization or other types of disinfection are required to disinfect the slurry that goes out of the anaerobic digester. This process is however mainly applied before the anaerobic treatment to ensure that the biogas producing bacteria are not deceased or disturbed in their processing.
In addition does the feedstock need to be heated to the optimal temperature before reaching the digester and thus a hygienization in the pre-treatment is the preferable approach to maximize the heat recovery (Ariunbaatar et al. 2014; Jørgensen 2015; Fløan 2015).
Pasteurization can be accomplished at many different temperatures and thus the HRT varies.
The main alternatives are 52C/ 10 h, 53.5C/8h, 55C/6h and 70C/1h (Ariunbaatar et al. 2014).
Cambi has developed another technology that uses 150C/20min and they claim that the methane yield and degrability increases for the feedstock, this is in particular good for cellulose containing materials such as grass, garden wastes and horse manure (Fløan 2015). In Ariunbaatar et al. (2014)’s study has they indentified some benefits for several pretreatment tetchnologies. They indentified, in their literature study, that 70C/2 h and 150C/1 h pretreatment
3 Impact such as marine and freshwater eutrophication which mainly affect this compartment has been excluded in this study due to high levels of uncertainty. N and P leaching are however, implemented in the LCA model in SimaPro 8 for later studies and development.
18 temperatures resulted in mehane increases of 2.69% and 11.9%, respectiviely. This comparison is done for mesophilic continous flow treatmet and is relative to no pretreatment applied.
2.7. Anaerobic digestion
Anaerobic digestion can be applied in several ways that highly affect the outcome of the treatment process (Raadal & Morken 2008; Søheim et al. 2010). There is several technologies of achieving anaerobic digestion, but only three most common digesting methods is mentioned in this chapter.
One stage batch
One stage batch is a process in which a liquefied slurry is fed into a digester, and then the substrate degrades. The digester is then emptied before next batch of organic slurry is to undergo anaerobic digestion. This is known as a one-stage treatment because all processing and degradation occur at the same time in the digester. This treatment option is simple to use, but provide lower biogas yield and is most commonly applied in small-scale farming (Luostarinen et al. 2011).
One stage continuous flow
This method applies the same simple one stage digestion treatment, but instead of emptying and refilling is the substrate continuously filled in, as some of the digested substrate is extracted from the digester. Such a treatment reduces the energy consumption by having a constantly heated main body of slurry in the digester. In addition, does the bacteria culture get a much more diverse nutrient mix and not necessary to start a new bacteria culture each time new substrate is added to the digester (Luostarinen et al. 2011).
CSTR (multi- stage continuous stirred flow)
This is the most common anaerobic treatment technology at current date for big scale treatment plants. The process is similar to the one stage continuous flow, as it treats new substrate as continuously emptied for old (Luostarinen et al. 2011). The difference is that this treatment option digests the substrate in separate stages and that the pasteurization is separated from the digestion. To convert long hydrocarbons into CH4 is several bacteria species needed. The pre- treatment processes can be size reduction of organic matter, hygienization, pre-separation, sonication, enzyme addition among others4 (Luostarinen et al. 2011). After the pre-treatment is
4 For a more detailed description of pre-treatment technologies is Luostarinen et al. 2011 recommended reading, pg. 18.
19 the substrate filled in the digester where continuously stirred to maximise the contact between the volatile solids and the bacteria to enhance the biogas production even more.
The last stage in a multi stage digester is the post treatment process. possible to recover a significant amount of the biogas emitted from a biogas plant by providing a cover or collection system. 10 – 30% of the biogas potential remain in the bioresidual, which in some cases can be further collected (Luostarinen et al. 2011). If not, will the bioresidual be emitted large amounts of methane to the atmosphere and thus contribute to GHG emissions as earlier described, (Luostarinen et al. 2011; Amon et al. 2006; Khalid et al. 2011; Muha et al. 2014). also possible to do a mechanical separation of the bioresidual where the wet and dry separation fractions are separated5.
2.8. Incineration
Incineration is the preferred solution for waste treatment in Norway at current date. In many cases is incineration applied for environmental reasons as incineration drastically reduces the need for landfill areas and thus the methane to air generation (Beylot & Villeneuve 2013).
Landfill often leads to leaching to groundwater and nearby rivers of organic compounds resulting in chemical oxidant formation, eutrophication etc. also an economical solution as some of the energy can be harvested from the combustion as both electricity and heat, where the latter is the most common appliance in Norway (Jørgensen 2015). By burning the organic material by itself it avoids sources for contamination and can therefore be used as filling in road construction or concreate production if the contaminations meets the quality requirements.
However only the bottom ash can be utilized for such a purpose. The incineration process creates two types of ash during combustion, fly ash and bottom ash. The fly ash is treated trough the off gas treatment system and can contain high levels of heavy metals and other cardiogenic compounds. Therefore, is the fly ash not suitable as a fertilizer and should be treated as special waste and landfilled in closed storages. The bottom ash leaves the heavier compounds that does not levitate during combustion and must therefore go through a sorting and later for heavy metal treatment and cleaning. therefore only “pure” ash with quality zero that can be used as a fertilizer. This can be achieved by restricting the inputs to only organic wastes where sewage sludge is considered a source of contamination and should therefore be limited in amount, (Boesch et al. 2014). There is several incineration types that are commercially used today. The two most common plant types is fluid gas bed and incineration moving grate (Jørgensen 2015).
5 For a more detailed description of the post-treatment technologies is Luostarinen et al. 2011 recommended reading, pg. 20.
